Environ Sci Pollut Res DOI 10.1007/s11356-014-2941-5

RESEARCH ARTICLE

Cyprodinil retention on mixtures of soil and solid wastes from wineries. Effects of waste dose and ageing Isabel Rodríguez-Salgado & Marcos Paradelo-Pérez & Paula Pérez-Rodríguez & Laura Cutillas-Barreiro & David Fernández-Calviño & Juan Carlos Nóvoa-Muñoz & Manuel Arias-Estévez

Received: 8 October 2013 / Accepted: 20 April 2014 # Springer-Verlag Berlin Heidelberg 2014

Abstract In spite of its wide-world economic relevance, wine production generates a huge amount of waste that threatens the environment. A batch experiment was designed to assess the effect of the amendment of an agricultural soil with two winery wastes (perlite and bentonite wastes) in the immobilization of cyprodinil. Waste addition (0, 10, 20, 40, and 80 Mg ha−1) and different times of incubation of soil-waste mixtures (1, 30, and 120 days) were tested. The addition of wastes improved the soil’s ability to immobilize cyprodinil, which was significantly correlated to total C content in soilwaste mixtures. Longer incubation times decreased the cyprodinil sorption possibly due to the mineralization of organic matter but also as a consequence of the high pH values reached after bentonite waste addition (up to 10.0). Cyprodinil desorption increased as the amount of waste added to soil, and the incubation time increased. The use of these winery wastes contributes to a more sustainable agriculture preventing fungicide mobilization to groundwater. Keywords Fungicides . Cyprodinil immobilization . Agricultural soils . Winery wastes . Bentonite . Perlite Responsible editor: Zhihong Xu Electronic supplementary material The online version of this article (doi:10.1007/s11356-014-2941-5) contains supplementary material, which is available to authorized users. I. Rodríguez-Salgado : M. Paradelo-Pérez : P. Pérez-Rodríguez : L. Cutillas-Barreiro : D. Fernández-Calviño : J. C. Nóvoa-Muñoz (*) : M. Arias-Estévez Área de Edafoloxía e Química Agrícola, Departamento de Bioloxía Vexetal e Ciencia do Solo, Facultade de Ciencias, Universidade de Vigo, 32004 Ourense, Spain e-mail: [email protected] M. Paradelo-Pérez Department of Agroecology, Faculty of Science and Technology, Aarhus University, Blichers Allé 20, P.O. Box 50, DK-8830 Tjele, Denmark

Introduction The vast use of agrochemicals in vineyards implies a high contamination risk of soils and groundwater, becoming a serious problem for agriculture and the environment (Komárek et al. 2010). “New-generation” organic-based fungicides with improved efficacy have been used from the past decade against fungal diseases which affect the quality and quantity of grape production (Arias et al. 2005). Cyprodinil (4-cyclopropyl-6-methyl-Nphenylpyrimidine) is a systemic fungicide recommended for the prevention and treatment of various rots of fungal origin that can affect fruit plants and vines. As a consequence, cyprodinil concentrations up to 462 μg kg−1 were found in vineyard soils (Bermúdez-Couso et al. 2007). Fortunately, soil organic matter possesses a positive influence on the immobilization of cyprodinil by soils (Arias et al. 2005; Pose-Juan et al. 2011), suggesting that the addition of organic amendments to agricultural soils could reduce the mobilization of this fungicide, diminishing the risk of surface and subsurface water contamination. In general, agrofood wastes are widely applied to agricultural soils as organic amendments to restore soil fertility and improve soil quality, being used in some cases as supplements or even substitutes of fertilizers (Diacono and Montemurro 2010; Golabi et al. 2007; Kulkarni et al. 2007; Madejón et al. 2001). Furthermore, recent studies demonstrated their utility in pesticide immobilization and degradation (Briceño et al. 2007; Fernández-Bayo et al. 2009; Marín-Benito et al. 2009). Traditionally, winemaking is a relevant agroindustry in several European countries in economic terms, mainly in France, Italy, and Spain. During the twentieth century, other countries have become important wine producers such as South Africa, Australia, USA, Chile, Argentina, Australia, and more recently, China and India. However, coupled with the global economic relevance of the wine sector, environmental issues related to the management of the solid and liquid

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wastes generated during winemaking have arisen. Following Martínez-Sabater et al. (2009), the world level of wastes derived from winemaking processes was estimated to be 2 million Mg years−1 of grape stalk, 2.3 million Mg years−1 of wine lees, 5.3 million Mg years−1 of grape marc, and 162 million m3 years−1 of winey wastewaters and vinasse. Due to this huge amount of waste produced and its potential impact on the environment, several alternatives for its reutilization have been assessed (Arvanitoyannis et al. 2006; Devesa-Rey et al. 2011; Ruggieri et al. 2009). Thus, although winery wastes were mainly used as amendments in agricultural soils due to notable contents of organic matter and plant nutrients (K, P, Ca, etc.), they were also successfully used in environmental issues as in the immobilization of heavy metals from contaminated waters (Farinella et al. 2008; Villaescusa et al. 2004) or pesticide retention in agricultural soils (Andrades et al. 2004; Fernández-Bayo et al. 2007; Pateiro-Moure et al. 2009). Filtration and clarification processes during wine elaboration in wineries generate a considerable amount of waste. Thus, expanded perlite is an alumina silicate of volcanic origin that is used in grape must filtration, whereas bentonite is a mineral of the smectite group used for wine fining. During grape must filtration and wine fining, undesirable substances that negatively affect the quality of the final product are removed. As a result, two wastes occur, a perlite waste (WP) and a bentonite waste (WB). The high sorption capacity of perlite and bentonite is also conferred to their wastes, and thus, they show a notable enrichment in organic compounds such as carbon and nitrogen and inorganic ones such as phosphorous, potassium, and other substances from agrochemicals applied to the vineyards (Nóvoa-Muñoz et al. 2008). In spite of this, WP and WB have a considerably lower content of organic matter than other more traditional winery wastes such as grape stalks, grape marc, wine lees, or vinasses which are almost entirely organic. The suitability of perlite and bentonite wastes (WP and WB) as amendments in agricultural soils has already been assessed in previous studies (Arias-Estévez et al. 2007; Nóvoa-Muñoz et al. 2008). However, the effects of the addition of WP and WB to agricultural soils in the pesticide behavior have not yet been studied. Therefore, the main objective of this work was to evaluate how the addition of WP and WB to an agricultural soil affects to its capacity for pesticide (cyprodinil) retention. In addition, the influence of the waste dose and the changes of chemical properties of soil-waste mixtures with aging in cyprodinil retention will also be assessed.

Material and methods Fungicide cyprodinil Technical-grade cyprodinil was supplied by Riedel-de Häen (Seelze-Hannover, Germany) with purity higher

than 90 %. Some of the relevant characteristics are listed in Table 1. Soil-waste mixtures Waste perlite (WP) and waste bentonite (WB) were supplied by Adegas Cunqueiro S.L. (Castrelo de Miño, Ourense, Spain). The wastes were air dried and passed through a 1-mm mesh sieve previous to use in the experiments. The soil selected for the experiment was from the first 20 cm of an acid Arenic Regosol developed from two-mica granites that which was recently changed from a forest use to an agricultural use (vineyard). Once sampled, the soil was air dried and sieved through a 2-mm mesh. The main properties of wastes and soil are listed in Table 2. Soil-waste mixtures (SWM) were prepared adding different waste doses equal to 0 (control) 10, 20, 40, and 80 Mg ha−1 assuming an effective soil depth of 20 cm and a bulk soil density of 1 kg dm−3. The SWM were incubated at field capacity (≈20 % w/w) for 1 day (t=0), 30 days (t=1), and 120 days (t=2). During the period of incubation, the moisture content of the mixtures was checked every 3 days, and the weight of water loss was compensated by adding distilled water. After incubation, the soil-waste mixtures were dried at 30 °C and passed through a 2-mm sieve in order to destroy newly formed soil aggregates. Soil-waste mixture’s pH was measured in 1:2.5 soil/solution suspensions or 0.1 M KCl using a pH meter with a combined glass electrode. Electrical conductivity (EC) was measured with a Crison 524 conductivimeter in filtered extracts obtained from 1:10 soil/water suspensions. Total organic carbon and nitrogen contents were determined by elemental analysis on a Thermo Finnigan 1112 series NC instrument (Waltham, MA, USA). Effective cation-exchange capacity Table 1 Chemical structure and properties of fungicide cyprodinil Chemical name (IUPAC)

4-Ciclopropil-6-metil-N-fenilpirimidin2-amina

Chemical structure

Substance group Molecular weight (g mol−1) Water solubility (g L−1) at 25 °C. Dissociation constant (pKa) at 25 °C Kow

Anilinopyrimidine 225.3 0.02 (pH=5), 0.013 (pH=7), and 0.015 (pH=9) 4.44 4.0

Environ Sci Pollut Res Table 2 Values of some chemicals parameters of soil and wastes (bentonite and perlite) used in this study

pH H2O pH KCl Sand Silt Clay C N C/N Naex Kex Caex Mgex Alex eCEC E.C.

% % % g kg−1 g kg−1 cmolc kg−1 cmolc kg−1 cmolc kg−1 cmolc kg−1 cmolc kg−1 cmolc kg−1 dS m−1

Soil

Bentonite waste

Perlite waste

5.3 4.1 68.5 20.3

5.1 5.3 – –

7.4 5.9 – –

11.2 0.4 0.2 2.0 1.6 0.03 0.14 0.08 1.5 3.4 0.1

– 164 19.3 8.5 2.6 164.9 0.68 1.55 0.8 170.5 7.0

– 122 18.2 6.7 2.6 27.7 0.27 0.62 0.1 31.3 0.6

The subscript “ex” refers to exchangeable cations eCEC effective cation-exchange capacity, E.C. electrical conductivity

(eCEC), i.e., CEC at soil pH, was estimated as the sum of exchangeable base cations (Cae, Mge, Nae, and Ke) extracted with 0.2 M NH4Cl and Al (Ale) displaced with 1 M KCl (Sumner and Miller 1996). Cations were measured by atomic absorption (Ca, Mg, Al) or flame emission (Na, K) spectrophotometry with a Thermo-Elemental Solar spectrometer. Sorption–desorption experiments Batch experiments were performed to investigate the effect of waste type, dose, and incubation time on the retention of cyprodinil. To this, 1 g of each soil-waste mixture was suspended in 10 mL of cyprodinil solutions ranging from 0.5 to 10 mg L−1, containing 0.01 M CaCl2 as a background electrolyte. The suspension was shaken end-over-end for 24 h at 25 °C in the dark. This shaking time was reported to be sufficient to reach the equilibrium between soil and fungicide solutions (Arias et al. 2005). After shaking, the suspensions were centrifuged 20 min at 2,000 rpm in a Rotina 35R centrifuge (Hettich Zentrifugen, Tuttlingen, Germany), and the resulting supernatant was taken and passed through a fiberglass filter with a 0.45-μm pore size from Macherey-Nagel (Düren, Germany). Previous tests revealed that these filters adsorb no cyprodinil. The concentration of cyprodinil in the supernatant was determined by HPLC-UV, and the amount of fungicide sorbed by the soil-waste mixtures was calculated by subtracting that present in solution after 24 h of equilibrium from the amount initially added. Adsorption tests were carried

out in duplicate. The results were modeled by fitting to the equations of Freundlich (Eq. (1)) and Langmuir (Eq. (2)): X ¼ K f Cn

X ¼

KLX mC 1 þ KLC

ð1Þ

ð2Þ

where X is the mass of adsorbed cyprodinil per unit mass of dry soil at equilibrium (mg kg−1); C is the concentration of cyprodinil in solution at equilibrium after 24 h (mg L−1), and Kf and n are the Freundlich coefficients. Kf (L kg−1) can be interpreted as the amount of sorbate adsorbed at C=1, and n (dimensionless) is a measure of heterogeneity in adsorption sites (Bermúdez-Couso et al. 2011). KL is a Langmuir constant associated with the energy of adsorption (L mg−1) and Xm is the maximum adsorption capacity of the soil samples (mg kg−1). The optimum parameter values for these equations were determined by non-linear regression analysis with SPSS software (IBM Corp., Armonk, NY, USA). For the desorption tests, immediately following adsorption of cyprodinil in soil-waste mixtures, the centrifuged residues were weighed to determine the amount of occluded solution and resuspended in 10 mL of 0.01 M CaCl2 solution containing no cyprodinil. The new suspensions were equilibrated for 24 h. Then, the samples were centrifuged and the supernatant filtered according the procedure explained before, and the amount of cyprodinil was then determined in the supernatant by high-performance liquid chromatography (HPLC). Desorbed cyprodinil was calculated by subtracting the occluded cyprodinil (estimated as the difference between the final and initial weights) from that measured in the supernatant after desorption test. In order to assess potential cyprodinil degradation during adsorption/desorption experiments, six selected soil-waste mixtures showing different aging and rate of waste addition were suspended in cyprodinil solutions as was described previously. Fungicide recovery was considered as the sum of cyprodinil concentration in the centrifuged supernatants and the amount of cyprodinil that remain in the soil, which was extracted following Rial-Otero et al. (2004). Fungicide analysis High-performance liquid chromatography (HPLC) analyses were performed using a liquid chromatograph (Dionex Corp. Sunnyvale, USA) equipped with a P680 quaternary pump, an ASI-100 autosampler, a TCC-100 Thermostatted Column Compartment, and a UVD170U detector and linked to a PC running the 6.8 version Chromeleon software program (Dionex Corp., Sunnyvale, USA). Separation was performed

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with a Luna C18 (150 mm×4.6 mm i.d., 5.0-μm particle size) analytical column obtained from Phenomenex (Madrid, Spain) and a guard column (4.0 mm×3.0 mm i.d., 5.0-μm particle size) containing the same packing material. The temperature of the HPLC column was kept constant at 40 °C. The mobile phase was acetonitrile and water with the following gradient: 55 % of acetonitrile was increased to 90 % in 7 min, held for 1 min, and then decreased to 55 % in 0.1 min, giving a total analysis time of 25 min after taking into account the equilibrium time. The injection volume was set to 50 μL at a HPLC flow rate of 1.0 mL min−1. Fungicide detection was carried out at 210 nm. The retention time in the chromatogram was 7 min.

Results and discussion Evolution of mixture properties during incubation The pHw of 1-day incubated mixtures remained constant with different doses of WB and WP ranging from 5.2 to 5.4 (Table 3). However, pHw increased for WB mixtures with longer incubation times, reaching values of 9.9 and 10.1 for WB 80 Mg ha−1 after 30 and 120 days, respectively. Incubation times of WP mixtures did not affect pHw. Changes in pHk presented the same behavior as pHw. However, pHk of WB mixtures incubated for 1 day increased from 4.1 for the control to 4.9 for 80 Mg ha−1 of WB. The increase of pHw for WB mixtures can be related with organic matter mineralization being more relevant for greater doses and longer incubation times. This is consistent with the perception of ammonia volatilization when mixtures were manipulated to keep the moisture constant during incubation. This suggests the ammonification of organic N from the wastes, which represents the highest nitrogen fraction (Arias-Estévez et al. 2007; Nóvoa-Muñoz et al. 2008). Other studies found a pH increase after adding winery wastes to soil (Bustamante et al. 2007) due to the mineralization of the organic matter. Electrical conductivity (EC) in the SWM increased with adding both residues, although WB drove greater EC values than WP (Table 3). However, incubation time decreased EC values in WB mixtures, from 842 μS cm−1 for 1 day to 504 and 540 μS cm−1 for 30 and 120 days, respectively. Bentonite and perlite wastes supplied C and N to soil, increasing their total contents in the mixtures in a proportional way, as for waste addition. Thus, the highest value of C content was 6.7 g kg−1 for 80 Mg ha−1 WB mixture; meanwhile, the C content was 4.9 for 80 Mg ha−1 WP mixture (Table 3). These differences were expected regarding the C content in the residues (Table 2). For both mixtures, C content decreased with the incubation time (Table 3) as a result of the mineralization of the organic

matter. Degradation was faster for WB than WP; C dropped almost 50 % for 80 Mg ha−1 WB mixture but only 5 % for 80 Mg ha−1 WP mixture after 30 days. The differences in organic content and composition led the different mineralization dynamics for each waste. Total N contents (Table 3) ranged from 0 to 0.9 g kg−1 and followed the behavior of total C contents due their relation with mineralization processes and as ammonia losses. Organic matter and nutrient levels give us an idea of the s u i t a b i l i t y o f t h e r e s i d ue s f or a gr on o m ic re us e (Arvanitoyannis et al. 2006; Devesa-Rey et al. 2011). The C and N losses from waste-soil mixtures with time suggested the mineralization of the organic matter and the subsequent decrease of nutrients levels. Also, organic matter humification can improve the sorption capacity of contaminants by the soilwaste mixtures. Faster C losses for WB- than WP-amended soil samples can be caused by differences in organic C composition; the presence of labile C favors the mineralization speed (Bustamante et al. 2007). Nitrogen losses can be related to the coarse texture of the soil that provides a better aeration and, therefore, favors the N volatilization (Hernández et al. 2002). Potassium was the dominant cation in the cationic exchange complex of the residue soil mixture and ranged from 0.1 cmolc kg−1 for control to 9.5 cmolc kg−1 for 80 Mg ha−1 WB mixture and 120 days (Table 3). Exchangeable K levels were greater for WB mixtures than WP mixtures as expected regarding the K content of the residues (Table 2). Incubation time increased the Kex levels for every mixture, especially from 30 to 120 days. Other cations in the effective cation-exchange capacity (eCEC) followed the sequence Na>Ca>Mg regarding their abundance, and their levels tend to diminish with time (Table S1, Supporting Information). Exchangeable Al was higher in soil control due to the acidic character of the soil used and decreased with the addition of residue by the increase of pH. Effective cation-exchange capacity (eCEC) increased with the addition of residue but, contrary to Kex evolution, decreased with time in all cases except for 80 Mg ha−1 WB. The addition of residues promoted the increase of eCEC levels and is in agreement with previous works (Arias-Estévez et al. 2007; Nóvoa-Muñoz et al. 2008). Organic matter from residues increased the number of exchange sites, but its degradation decreased the exchange sites after incubation (Croker et al. 2005). This increase in eCEC should improve the retention capacity of contaminants in the amended soil (NóvoaMuñoz et al. 2008), which is especially notorious for coarse texture soils such as the soil used in this study. Sorption studies Sorption curves of cyprodinil by both soil-residue mixtures (for four residue doses and three incubation times) are plotted

0.0 (0.0) 0.4 (0.0) 0.4 (0.0) 0.5 (0.0) 0.9 (0.0) 0.3 (0.0) 0.3 (0.0) 0.5 (0.0) 0.8 (0.0)

0.0 (0.0) 0.2 (0.1) 0.2 (0.1) 0.3 (0.0) 0.4 (0.0) 0.1 (0.0) 0.2 (0.0) 0.4 (0.0) 0.7 (0.0)

0.0 (0.0) 0.1 (0.0) 0.2 (0.0) 0.3 (0.0) 0.4 (0.0) 0.1 (0.0) 0.2 (0.0) 0.3 (0.0) 0.7 (0.0)

120

4.2 (0.1) 4.2 (0.0) 5.0 (0.5) 7.1 (0.1) 9.2 (0.0) 4.1 (0.0) 4.1 (0.0) 4.0 (0.1) 4.1 (0.0)

4.2 (0.2) 4.4 (0.1) 4.7 (0.2) 6.1 (0.7) 9.7 (0.4) 4.3 (0.2) 4.2 (0.2) 4.2 (0.1) 4.4 (0.2)

WB waste bentonite, WP waste perlite

0.1 (0.1) 0.9 (0.0) 1.5 (0.1) 3.7 (0.4) 5.3 (1.6) 0.3 (0.0) 0.4 (0.0) 0.7 (0.4) 1.1 (0.0)

1

30

4.1 (0.0) 4.0 (0.0) 4.2 (0.0) 4.6 (0.0) 4.9 (0.0) 4.1 (0.0) 4.0 (0.0) 4.0 (0.0) 4.1 (0.0)

1

5.4 (0.2) 5.8 (0.1) 6.5 (0.2) 7.3 (0.6) 10.1 (0.1) 5.7 (0.3) 5.5 (0.1) 5.3 (0.5) 5.3 (0.2)

120

Kex (cmolc kg−1)

5.5 (0.4) 6.2 (0.1) 6.7 (0.5) 8.9 (0.1) 9.9 (0.0) 5.4 (0.2) 5.7 (0.6) 5.0 (0.5) 5.5 (0.1)

30

N (g kg−1)

5.3 (0.1) 5.3 (0.0) 5.2 (0.0) 5.2 (0.0) 5.3 (0.0) 5.4 (0.0) 5.3 (0.1) 5.3 (0.0) 5.3 (0.0)

1

120

1

30

pHk

pHw

Numbers 10, 20, 40, and 80 represents the dose of residue in megagrams per hectare

Control WB-10 WB-20 WB-40 WB-80 WP-10 WP-20 WP-40 WP-80

Incubation (days)

Control WB-10 WB-20 WB-40 WB-80 WP-10 WP-20 WP-40 WP-80

Incubation (days)

0.1 (0.0) 0.9 (0.0) 1.8 (0.4) 2.8 (0.1) 6.5 (0.3) 0.2 (0.0) 0.4 (0.0) 0.7 (0.0) 1.1 (0.1)

30

82.3 (6.5) 163.5 (11.2) 232.3 (9.1) 386.3 (6.0) 841.5 (7.8) 138.9 (1.1) 157.2 (7.0) 101.2 (1.4) 126.3 (8.1)

1

E.C. (μS cm−1)

0.1 (0.0) 1.2 (0.1) 2.1 (0.1) 5.0 (0.3) 9.5 (0.8) 0.4 (0.0) 0.5 (0.1) 1.0 (0.1) 2.0 (0.3)

120

72.6 (3.5) 77.7 (3.8) 100.4 (6.3) 150.0 (5.2) 504.0 (2.8) 87.4 (8.5) 79.2 (3.6) 95.2 (4.2) 137.1 (17.0)

30

5.5 (0.5) 6.1 (0.5) 6.6 (0.3) 8.5 (0.7) 10.4 (1.4) 5.1 (0.4) 5.2 (0.2) 5.4 (0.7) 5.7 (0.3)

1

0.4 (0.1) 1.0 (0.1) 1.8 (0.3) 2.3 (0.1) 3.5 (0.1) 1.0 (0.1) 1.6 (0.1) 2.8 (0.1) 4.7 (0.2)

30

4.9 (0.5) 4.7 (0.2) 4.7 (0.7) 5.8 (0.1) 9.8 (0.0) 4.9 (0.1) 4.6 (0.3) 4.8 (0.2) 4.9 (0.3)

30

0.3 (0.0) 1.1 (0.0) 1.9 (0.0) 3.5 (0.1) 6.7 (0.2) 0.9 (0.0) 1.5 (0.1) 2.6 (0.1) 4.9 (0.1)

1

eCEC (cmolc kg−1)

88.0 (11.9) 92.1 (12.6) 111.7 (9.4) 227.4 (36.7) 540.0 (9.1) 93.3 (6.1) 91.7 (6.2) 102.5 (8.7) 140.9 (5.9)

120

C (g kg−1)

Table 3 Mean values and standard deviation (between brackets) of some chemical properties of the soil-waste mixtures after different incubation times (1, 30, and 120 days)

4.3 (0.4) 4.2 (0.5) 4.1 (0.5) 7.7 (0.9) 11.0 (1.2) 4.7 (0.4) 4.5 (0.7) 4.6 (0.5) 4.6 (1.1)

120

0.4 (0.0) 0.9 (0.1) 1.5 (0.1) 2.0 (0.0) 3.1 (0.1) 0.9 (0.0) 1.5 (0.1) 2.5 (0.0) 4.6 (0.2)

120

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in Fig. 1. In general, adsorption curves for shorter incubation times were C-type linear curves. With longer times, i.e., as a result of aging of SWM, retention curves were close to L-type, suggesting progressive saturation of the soil (Limousin et al. 2007). Other authors found similar curves for different pesticides such as metalaxyl or penconazole (Arias et al. 2006; Marín-Benito et al. 2012a). Retention curves of cyprodinil were successfully fitted to the Freundlich model (Eq. 1) showing r2 values above 0.95 (Table 4). Kf and n coefficients are shown in Table 4. For t=0, Kf increased linearly with the residue dose, from 24.4± 0.9 L kg−1 for control to 73.9±3.9 L kg−1 for WB-80 and 42.3±1.5 L kg−1 for WP-80. WB mixtures almost doubled the WP retention capacity.

30 days

1 day

Waste Bentonite

120 days

Fig. 1 Sorption curves of cyprodinil by SWM depending on dose and type of waste added and the incubation time. WB waste bentonite, WP waste perlite; 10, 20, 40, and 80 Mg ha−1 are the doses of residue added to soil. a, c, and e the cyprodinil sorption curves of WB mixtures for 1, 30, and 120 incubation days, respectively. b, d, and f the cyprodinil sorption curves of WP mixtures for 1, 30, and 120 incubation days, respectively

We studied the relation between the cyprodinil sorption capacity, represented by Kf values, and the properties of the SWM. Values of Kf and total C content were significantly correlated (n=9, r=0.872, pWP-10>WP-20>WP-40>WP-80> WB-10>WB-20>WB-40>WB-80. In control samples, i.e., in soil samples without waste addition, cyprodinil desorption tended to decrease with the concentration of fungicide previously added. Thus, desorption percentages vary from 35 to 26 % as a response to previous additions of 0.5 and 10 mg L−1 of cyprodinil, respectively. A similar trend was observed in WP-amended soils in which the percentages of cyprodinil desorption, for fungicide additions of 0.5 and 10 mg L−1, diminished between 3 and 8 % (34 to 26 % for WP-10, 26 to 22 % for WP-20, 28 to 23 % for WP-40, and 21 to 18 % for WP-80). On the contrary, the desorbed percentage of cyprodinil in soils amended with WB increased as the fungicide addition increased, achieving maximum values between 16 and 26 % depending on the amount of WB applied to the soil in response to the addition of 10 mg L−1 of cyprodinil (Table 5). Aging of the soil-waste mixtures, as a consequence of the increase in the incubation time, resulted in greater percentages of cyprodinil desorption, especially in SWM-amended soils at the highest rates of WB (40 and 80 Mg ha−1). Thus, WB-40amended soil samples showed the highest rates for cyprodinil desorption after 30 days of incubation with values in the range 18 to 50 %, being slightly lower for 120 days of incubation, ranging from 13 to 27 % (Table 5). This behavior in the cyprodinil desorption followed the pH evolution for this SWM (Table 1), which presented a maximum after 30 days of incubation. In the case of the soil samples amended with 80 Mg ha−1 of WB, cyprodinil desorption increased monotonically with the incubation time from average of 10 % (1 day) to 29 % (30 days) and 35 % (120 days). This general increase of cyprodinil desorption with time suggested a higher mobility of the pesticide in more aged mixtures which could be related to a greater mineralization of organic matter and an increase of pH, ultimately leading to a higher mobility of cyprodinil. However, in spite of the role of mineralization of organic matter in the potential mobilization of cyprodinil, the addition of organic matter to agricultural soil has been reported as a practice to ensure the immobilization of this pesticide (Arias et al. 2005).

Conclusions In conclusion, both winery wastes increased the soil capacity to immobilize cyprodinil, although it was somewhat greater in those samples amended with bentonite waste (WB) than with perlite waste (WP). The cyprodinil sorption seems to be controlled by the organic C content in the soil-waste mixtures, showing differences between both tested winery wastes, but decreasing with the mineralization of soil organic matter during aging (i.e., as the incubation time increases). In addition, the increase in the pH of soil-waste mixtures as a consequence of the addition of WB could influence the cyprodinil sorption with regard to competing to sorption sites against dissolved organic C. Mobilization of previously sorbed cyprodinil, i.e., desorption, increases as the amount of added fungicide increased, as well as with the aging of the soil-waste mixtures due to mineralization of the soil organic matter. In terms of agricultural management, the use of these winery wastes could contribute to a more sustainable agriculture, preventing pesticide mobilization to groundwater. However, a greater efficiency in the cyprodinil immobilization will be expected if waste addition to soil takes place close to the fungicide application. Acknowledgments This work was supported by Consellería de Medio Rural from Xunta de Galicia (FEADER2009-22) through the CO-106-09 contract between the Soil Science and Agricultural Chemistry Area of the University of Vigo and Bodegas Cunqueiro. The financial support of CIA3 through FEDER founds under the program of Consolidation and Arrangement of Research Units from Consellería de Educación (Xunta de Galicia) is also acknowledged. M. Paradelo is supported by a postdoctoral grant from the Barrié de la Maza Foundation and P. Pérez-Rodríguez is funded by the FPU program from the Spanish Ministry of Education.

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